Marine Pollution Bulletin 45 (2002) 3–16
www.elsevier.com/locate/marpolbul
Keynote papers
Cell bioassays for detection of aryl hydrocarbon (AhR) and
estrogen receptor (ER) mediated activity in environmental samples
J.P. Giesy
a
a,*
, K. Hilscherova a, P.D. Jones a, K. Kannan a, M. Machala
b
Department of Zoology, National Food Safety and Toxicology Center, Institute for Environmental Toxicology, Michigan State University,
East Lansing, MI 48824, USA
b
Veterinary Research Institute, Hudcova 70, 621 32 Brno, Czech Republic
Abstract
In vitro cell bioassays are useful techniques for the determination of receptor-mediated activities in environmental samples
containing complex mixtures of contaminants. The cell bioassays determine contamination by pollutants that act through specific
modes of action. This article presents strategies for the evaluation of aryl hydrocarbon receptor (hereafter referred as dioxin-like) or
estrogen receptor mediated activities of potential endocrine disrupting compounds in complex environmental mixtures. Extracts
from various types of environmental or food matrices can be tested by this technique to evaluate their 2,3,7,8-tetrachlorodibenzop-dioxin equivalents or estrogenic equivalents and to identify contaminated samples that need further investigation using resourceintensive instrumental analyses. Fractionation of sample extracts exhibiting significant activities, and subsequent reanalysis with the
bioassays can identify important classes of contaminants that are responsible for the observed activity. Effect-directed chemical
analysis is performed only for the active fractions to determine the responsible compounds. Potency-balance estimates of all major
compounds contributing to the observed effects can be calculated to determine if all of the activity has been identified, and to assess
the potential for interactions such as synergism or antagonism among contaminants present in the complex mixtures. The bioassay
approach is an efficient (fast and cost effective) screening system to identify the samples of interest and to provide basic information
for further analysis and risk evaluation.
Ó 2002 Elsevier Science Ltd. All rights reserved.
Keywords: In vitro cell bioassays; Dioxin-like activity; Estrogen receptor-mediated activity; Complex mixtures; Fractionation; Toxic equivalents;
Endocrine disruptors
1. Introduction
There is increasing concern over the potential adverse
effects of xenobiotics present in the environment and
foodstuffs on human and wildlife populations. Two
groups of toxicants of current interest are dioxin-like
and (anti) estrogenic chemicals. Many of these ubiquitous compounds are hydrophobic, lipophilic and resistant to biological and chemical degradation. These
properties impart persistency and propensity to bioaccumulate and biomagnify to concentrations that can
cause deleterious effects on cells and tissues. In the environment, chemicals occur as complex mixtures including different congeners and isomers of both natural
and anthropogenic origin. The concentrations and toxic
*
Corresponding author.
E-mail address: jgiesy@aol.com (J.P. Giesy).
potencies of compounds present in complex mixtures
can range over several orders of magnitude. In addition,
interactions among different classes of compounds (e.g.,
estrogenic vs. anti-estrogenic) can modulate the toxic
potential. This complicates hazard evaluation and risk
assessment of complex mixtures of xenobiotics. Furthermore, toxic effects of some contaminants, even
those, which are analytically determined, are not well
characterized. There are many potentially significant
classes of contaminants that are not studied in detail,
primarily due to a lack of suitable instrumental techniques or analytical standards. In other words, chemical
analysis has been used to identify and quantify only
those chemicals for which analytical techniques and
standards are available. Instrumental analyses do not
account for interactions among the chemicals in complex mixtures and provide little information on their
biological effects. Chemical analyses can also be costly
0025-326X/02/$ - see front matter Ó 2002 Elsevier Science Ltd. All rights reserved.
PII: S 0 0 2 5 - 3 2 6 X ( 0 2 ) 0 0 0 9 7 - 8
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J.P. Giesy et al. / Marine Pollution Bulletin 45 (2002) 3–16
and time consuming. Thus, chemical analyses can
underestimate the potential risks posed by these chemicals; some toxicologically important compounds could
be overlooked.
In vitro cell bioassays offer a rapid, sensitive and
relatively inexpensive solution to some of the limitations
of instrumental analysis. They enable estimation of total
biological activity of all compounds that act through the
same mode of action present in extracts of any environmental media. Bioassays also integrate possible
interactions among chemicals. In this review, the applicability of in vitro cell bioassays for assessment of two
toxicological modes of action, dioxin-like toxicity and
estrogen receptor(ER)-mediated activity, is evaluated.
Several reviews concerning dioxin-like and estrogenic
activities of xenobiotics have appeared (Gray et al.,
1997; Gillesby and Zacharewski, 1998; Ankley et al.,
1998; van den Berg et al., 1998). In our paper, the
strategy of the cell bioassay approach for evaluation of
receptor mediated activity of complex mixtures is presented, including fractionation procedures, potency
balance calculations, toxicant identification and risk
assessment. Also, the classes of aryl hydrocarbon receptor (AhR)-agonists and compounds that have been
shown to elicit endocrine disrupting potential are summarized.
shock proteins dissociate from the complex and it forms
a dimer with the Ah receptor nuclear traslocator (ARNT)
protein and possibly other factors. The heteromeric ligand:AhR:ARNT complex binds with high affinity to
specific DNA sequences, the dioxin-responsive element
(DRE). The binding to the DRE results in DNA bending, disruption of chromatin and nucleosome and thus
increased promoter accessibility and transcriptional activation of adjacent responsive genes (see Fig. 1) (Denison and Heath-Pagliuso, 1998; Hankinson, 1995).
The traditional well-known ligands for AhR have
been described as hydrophobic aromatic compounds
2. Dioxin-like activity
Chemicals that elicit toxic effects similar to that of
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), known as
dioxin-like chemicals, are of great concern due to their
ability to cause hepatotoxicity, embryotoxicity, teratogenicity, immunotoxicity, dermal toxicity, lethality, carcinogenesis, wasting syndrome and tumor promotion in
many different species at low concentrations (Ahlborg
et al., 1992; Peterson et al., 1993). A number of studies
have demonstrated that several toxic and biochemical
effects caused by dioxin-like chemicals are mediated
through AhR (Nebert et al., 1993; Lucier et al., 1993).
The AhR, which belongs to the basic helix-loop-helix
protein family (Nie et al., 2001), is a ligand-dependent
transcription factor located in the cytosol, complexed
with heat shock proteins. It has been shown that the
strength with which congeners bind to the AhR is directly proportional to the toxicity, enhanced gene transcription and enzyme activities mediated by the AhR
mechanism (Safe, 1995b). The role of AhR in mediating
toxic and biological effects of dioxin-like chemicals has
been well documented in a number of studies, even
though the exact biochemical mechanism leading to the
wide spectrum of toxic responses is yet to be elucidated
(Denison and Heath-Pagliuso, 1998). After binding of
ligand to cytosolic AhR, the receptor ligand complex is
activated and translocated to the nucleus, where heat
Fig. 1. Mechanism of AhR- or ER- receptor-mediated response in cell
(adapted from Blankenship et al., 2000; Villeneuve et al., 1998). For
description see text. HSP ¼ heat shock proteins, P ¼ phosphates:
phospohorylation is an important regulatory factor for receptor
function.
J.P. Giesy et al. / Marine Pollution Bulletin 45 (2002) 3–16
with planar structure of a particular size, which fits the
binding sites (Poland and Knutson, 1982; Lewis et al.,
1986). Thus, the ability of these ligands to bind to the
AhR and to cause toxic effects depends on their structure and substitution pattern. These include planar
congeners of polychlorinated dibenzo-p-dioxins and
dibenzofurans (PCDDs and PCDFs), chlorinated azobenzenes and azoxybenzenes, polychlorinated biphenyls (PCBs), several polycyclic aromatic hydrocarbons
(PAHs) and polychlorinated naphthalenes (Blankenship
et al., 2000). Other chemicals suggested as potential
AhR agonists due to their stereochemical configuration,
but not yet experimentally confirmed, include polybrominated and chloro-/bromo-analogs of the previously
listed classes of compounds (Till et al., 1997), alkylatedchlorinated dioxins and furans, chlorinated dibenzothiophenes, chlorinated xanthenes and xanthones (Van
Den Heuvel et al., 1994), polychlorinated diphenyltoluenes, anisols, anthracenes, fluorenes and others (Sanderson and Giesy, 1998). New types of relatively weak
AhR ligands or inducers (compared to TCDD) have
been identified, which include both natural and synthetic compounds (Denison and Heath-Pagliuso, 1998).
These compounds deviate from the traditional criteria
of planarity, aromaticity and hydrophobicity. The natural compounds that bind to the AhR include, among
others, indoles, tryptophan-derived products, oxidized
carotinoids and heterocyclic amines. Some pesticides or
drugs with various structures, such as imidazols and
pyridines also possess AhR binding affinity. These ligands act as transient inducers and bind to the AhR
with weak affinity and are rapidly degraded by the induced detoxification enzymes.
3. Estrogenic activity
There has been increasing interest in chemicals that
can modulate the endocrine system. Such compounds
have the potential to disrupt normal reproduction or
developmental processes which can lead to adverse
health effects such as compromised reproductive capacity, breast and testicular cancer, reproductive dysfunction such as feminization or demasculanization of
males and other adverse effects. A wide range of compounds including natural products, pharmaceuticals and
industrial chemicals have been shown to be estrogen
mimics. Some hormone-mimicking chemicals can elicit
multiple endocrine disrupting activities that are mediated by various mechanisms of action, some of them
may be active only during certain stages of development
(Sohoni and Soto, 1998). Their effects can be mediated
through receptor-mediated mechanisms (such as estrogen or androgen receptor), but some compounds can
disrupt hormone functions at different levels of the endocrine system, not directly interacting with the recep-
5
tor. Estrogen-like compounds exert effects by resembling
those of estrogen but these effects not mediated by the
ER (Gillesby and Zacharewski, 1998). Various modes of
actions have been reported, which include binding of
chemical to other nuclear receptors, which then interact
with an estrogen responsive element (ERE); acting
through other receptors and/or signal transduction
pathways; modulations of steroidogenesis and catabolism of active steroid hormones (Machala and Vondracek, 1998). Estrogenic compounds are characterized
by their ability to bind to and activate the ER, which is a
transcription factor belonging to the nuclear receptor
family. While there are structural similarities among
some compounds that are ER agonists, other ER-active
compounds do not share similar structures. Upon
binding of an estrogenic compound to the ligand binding domain of the ER (located predominantly in the
nucleus), the associated heat shock protein complex,
which masks the DNA binding domain, dissociates and
subsequently the ligand occupied receptor dimerizes.
The homodimer complex interacts with specific DNA
sequences referred to as EREs located in the regulatory
regions of estrogen-inducible genes. ER complexes
bound to an ERE recruit additional transcription factors, leading to increased gene transcription and synthesis of proteins required for expression of hormonal
action (Fig. 1) (Joyeux et al., 1997; Fielden et al., 1997).
A series of natural and synthetic endocrine disrupting
compounds have been identified by different in vivo and/
or in vitro methods. Numerous specific testing systems
have been developed for the detection of effects at different levels of the endocrine system (Villeneuve et al.,
1998; Gray et al., 1997). Examples of xenoestrogenic
compounds including natural and major classes of industrial contaminants are presented along with the
method used to determine their relative estrogenic potency (Tables 1 and 2).
4. Cell bioassay approaches
Bioassays based on the responses of either wild type or
genetically engineered eukaryotic cells enable the assessment of potencies of individual chemicals or complex
mixtures of environmental contaminants in extracts to
cause AhR- or ER-mediated effects. Either endogenous
responses or exogenous reporter systems, incorporated
into the cell, are used for the measurements. The induction of transcription by the responsive genes following
the exposure of cells to specific ligands or mixtures of
compounds can be assessed by measuring endogenous or
genetically engineered responses such as protein expression by measuring the amount of protein directly or by
measuring an enzyme activity.
The endogenous responses for AhR binding, increased
expression and induced activity of cytochrome P4501A1
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Table 1
Examples of endocrine disrupting compounds: natural products
Compound
Mode of action
Assay
Reference
Phytoestrogens
Indole-3-carbinol
ER agonist
RER (MCF-7-luc), YES
ER agonist, androgenic
after metabolization
ER agonist
YES, in vivo fish
RER (MCF-7-luc), YES
Decreased aromatase
enzyme activity
In vitro ER mediated
PAP induction
In vitro human cell
culture system
Villeneuve et al. (1998), Gaido
et al. (1997)
Gaido et al. (1997), StahlschmidtAllner et al. (1997)
Villeneuve et al. (1998), Gaido
et al. (1997)
Markiewicz et al. (1993)
ER agonist
RER (ER-CALUX)
Estrogenic
In vitro and in vivo
vitellogenin production
In vitro and in vivo
vitellogenin production
In vitro ER mediated
PAP induction
CB-ER, RER, RER
(MVLN)
b-Sitosterol
Coumestrol
Enterolactone, enterodiol
Bioflavanoids
Genistein
Biochanin A, daidzein, equol
Quercetin, naringenin, luteolin apigenin, chrysin,
kaempferol, hydroxy- and methoxy-flavones
Mycoestrogens
Zearalenone
ER agonists, estrogenic
Estrogenic, antiestrogenic, ER agonists
ER agonist
CB-ER, RER, VTG
in vitro
Wang et al. (1994)
Markiewicz et al. (1993),
Legler et al. (1998)
Nimrod and Benson (1997)
Nimrod and Benson (1997)
Markiewicz et al. (1993)
Kuiper et al. (1998), Safe Gaido
(1998), Le Bail et al. (1998)
Kuiper et al. (1998),
Celius et al. (1999)
YES, yeast based recombinant ER-reporter assay; E-screen, MCF-7 cell proliferation; CB-ER, in vitro competitive receptor binding assay; RER,
in vitro recombinant receptor-reporter cell bioassay; VTG-in vitro, in vitro vitellogenin synthesis in cultured male trout hepatocytes.
monooxygenase activities, such as 7-ethoxyresorufin
O-deethylase (EROD) or aryl hydrocarbon hydroxylase
are measured as markers of responses to AhR agonists
(Tillitt et al., 1991; Hahn et al., 1996). ER-mediated activity can be examined by the determination of specific
gene products such as vitellogenin, pS2 or steroid hormone binding globulins (Villeneuve et al., 1998; Sumpter
and Jobling, 1995; Pelissero et al., 1993). Some animal
and human cell lines used for the detection of in vitro
TCDD-like or estrogenic activity are listed (Table 3).
Recombinant cell lines are prepared by transient or
stable transfection of wild type cells with reporter genes
under transcriptional control of either DRE or ERE.
Transient transfection is relatively fast and easy to perform, but variations in transfection efficiency warrant
the need for co-transfection with internal constitutive
control to correct results for the transfection efficiency.
It can be used only for short-term studies, because
transgenes are usually lost after about 72 h. In addition,
physiological conditions including target DNA accessibility or overexpression of receptors or target DNA do
not reflect the normal cell function (Joyeux et al., 1997).
Stable transfection requires co-transfection of the plasmid with the gene of interest and plasmid encoding the
marker for drug resistance, enabling selection of only
successfully transfected cells and their development into
a stable cell line. The gene of interest becomes a permanent part of the cell genome. These cell lines are
suitable for longer-term experiments and their results
are more reproducible. Transfection with recombinant
expression vectors, which contain selected responsive
elements upstream of a reporter gene produces a cell
bioassay for specific class of chemicals. The most common reporter genes are firefly luciferase (luc), alkaline
phosphatase (PAP), chloramphenicol acetyl transferase,
or b-galactosidase (Joyeux et al., 1997; Zacharewski,
1997). Either native AhR or ER is used or the cells can
be co-transfected with a chimeric receptor and the recombinant reporter gene. Introduction of a complete
receptor–reporter system into the cells enables the development of responsive bioassays from cells with no
endogenous receptor present, such as in yeast cells
(Klotz et al., 1996). Recombinant yeast cells are easy to
develop and maintain and they are free of nuclear receptor background, which causes potential interference
in the assay. However, yeast cells are not necessarily a
good system to screen for effects in animal cells because
of differences in ligand specificities between animal cells
and yeast and the ability of yeast to metabolize proestrogens to estrogens (Villeneuve et al., 1998).
J.P. Giesy et al. / Marine Pollution Bulletin 45 (2002) 3–16
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Table 2
Examples of endocrine disrupting compounds: synthetic compounds
Compound
Mode of action
Assay
Reference
YES
in vitro cell line tests, in vivo
E-screen and other effects
Nafoxidine, clomiphene
Ethynylestradiol
Antiandrogenic activity
Antiestrogenic drug
binding to ER, antagonist or agonist
Antiestrogenic and antiandrogenic activity
ER agonist
ER agonist
Sohoni and Soto (1998)
Taylor et al. (1984), Ramkumar and Alder
(1995), Shelby et al. (1996), Favoni and Cupis
(1998)
Sohoni and Soto (1998), Favoni and Cupis
(1998)
Gaido et al. (1997)
Nimrod and Benson (1997), StahlschmidtAllner et al. (1997)
Additives
Parabens
ER agonists
t-Butylhydroxyanisol
Estrogenic
Pharmaceuticals
Flutamide
Tamoxifen
Hydroxytamoxifen
Pesticides
Insecticides
o,p0 -DDT
o,p0 -DDD, o,p0 -DDE
p,p0 -DDE
p,p0 -DDD
p,p0 -DDT
Kepone
ER agonist, antiandrogenic activity
ER agonists
Androgen receptor antagonist, weak ER and
androgen receptor
agonist
Antiadrogenic and weak
antiestrogenic activity
ER agonist
ER agonist, estrogenic
ER agonist,
estrogenic—after
metabolization
ER agonist
Endosulfan, Dieldrin,
Lindane
Toxaphene
Methyl parathion
Estrogenic
Estrogenic
Chlordecone
Chlordane
Estrogenic
ER agonist
Methoxychlor
ER agonist—after
metabolization
Endocrine modulators,
non-ligand binding
Carbamate insecticides
(Aldicarb, Bendiocarb,
Cabaryl, Methomyl,
Oxamyl)
Pyrethroid insecticides
(Sumithrin, Fenvalerate,
D -trans Allethrin,
Permethrin)
Fungicides
Vinclozolin
Dodemorph, Triadimefon, Biphenyl
Herbicides
Atrazine
Simazine
YES, E-screen and other effects
YES
In vitro, in vivo
CB-ER, YES, in vivo uterotropic
response
E-screen
Routledge et al. (1998)
Soto et al. (1995)
YES, RER (ER-CALUX),
VTG-in vitro
YES
CB-androgen receptor, in vivo mice
study
Sohoni and Soto (1998), Gaido et al. (1997),
Legler et al. (1998), Sumpter and Jobling (1995)
Gaido et al. (1997)
Kelce et al. (1995)
YES
Sohoni and Soto, 1998
YES, CB-ER, RER (MCF-7-luc)
E-screen
RER (ER-CALUX), E-screen,
in vitro þ in vivo
Klotz et al., 1996
Nimrod and Benson, 1997
Legler et al. (1998), Nimrod and Benson (1997),
Shelby et al. (1996)
RER (ER-CALUX)
Legler et al. (1998)
E-screen
YES, VTG—in vitro
In vivo effects on estrus cycle in
mice
YES, VTG—in vitro
RER (ER-CALUX)
In vivo—effects on endocrine
function in mice
RER (ER-CALUX) in vitro þ
in vivo
in vitro modulation of estrogen and
progesterone receptor in human
breast and endometrial cancer cells
Soto et al. (1994)
Petit et al. (1997)
Asmathbanu and Kaliwal (1997)
Petit et al. (1997)
Legler et al. (1998)
Cranmer et al. (1984)
Legler et al. (1998), Shelby et al. (1996)
Klotz et al. (1996)
Estrogenic (different
mechanisms)
In vitro pS2 gene expression
E-screen
Go et al. (1999)
Antiandrogen
Kelce et al. (1994), Sohoni and Soto (1998)
Estrogenic
In vitro androgen receptor binding
assay, YES
YES, VTG in vitro
Soto et al. (1994)
Estrogen, antiestrogen
Antiestrogen
RER (MCF-7-luc), in vivo
In vivo
Villeneuve et al. (1998)
Tennant et al. (1994)
(continued on next page)
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Table 2 (continued)
Compound
Mode of action
Assay
Reference
ER agonists
Androgenic
YES, CB-ER, RER (MCF-7-luc)
Imposex in snails, various in vivo
effects in gastropods
Klotz et al. (1996)
Stahlschmidt-Allner et al. (1997), Matthiessen
and Gibbs (1998)
ER agonist, antiandrogenic activity
In vitro þ in vivo, E-screen, YES
ER agonist
In vitro þ in vivo
Sohoni and Soto (1998), Stahlschmidt-Allner
et al. (1997), Soto et al. (1995), Jobling et al.
(1995)
Stahlschmidt-Allner et al. (1997), Jobling et al.
(1995)
ER agonist, estrogenic
RER (MCF-7-luc, ER-CALUX),
YES, number of in vitro and in vivo
assays, E-screen, Vtg-in vitro
Octylphenol
ER agonist
Butylphenol, Pentylphenol
Nonylphenol polyethoxylates and polyethoxycarboxylates
Pentachlorophenol
Estrogenic
RER (MCF-7-luc)
Number of in vitro and in vivo
assays
E-screen
Alachlor, Nonachlor
Tributyltins
Industrial chemicals
Phthalates
Butylbenzylpthalate
Dibutylpthalate
Alkylphenols
Nonylphenol
Bisphenol A
Number of in vitro and in vivo
assays
Servos (1999)
Decrease in blood testosterone concentration
ER agonist
In vivo ewes feeding study
Beard et al. (1999)
RER (MCF-7-luc, ER-CALUX),
YES, VTG in vitro
YES
Villeneuve et al. (1998), Gaido et al. (1997)
Persistent organic pollutants
PCDD
antiestrogenic—different
mechanisms
PCBs
ER agonists or antagonists or other mechanism—depending on the
substitution
Arochlor 1260 (PCBs
Estrogenic, effect on sexmixture), Arochlor 1260
ual differentiation, gonadal abnormalities
Hydroxy-PCBs
ER agonists or antagonists
6-hydroxy chrysene
Heavy metals
Cations of cadmium, cobalt, copper, mercury,
nickel, zinc
Cadmium
Lead
Abbreviation as in Table 1.
Nimrod and Benson (1997), Soto et al. (1995)
ER agonists
Antiandrogenic activity
PAHs
Villeneuve et al. (1998), Gaido et al. (1997),
Legler et al. (1998), Servos (1999), Nimrod and
Benson (1997), Shelby et al. (1996), Soto et al.
(1995)
Servos (1999)
Soto et al. (1995)
ER agonists—estrogenic,
antiestrogenic—different
mechanisms
Antiestrogenic
Depression or increases
in testosterone production
Decrease in plasma testosterone and cortisol
Modification of pituitary
hormone secretion
Delayed sexual maturation, suppression of sex
steroid biosynthesis
Legler et al. (1998), Soto et al. (1995)
Sohoni and Soto (1998)
In vivo þ in vitro studies
Safe and Krishnan (1995)
RER (transient MCF-7-luc),
E-screen, in vivo—vaginal cell
cornification in mice
Joyeux et al. (1997), Soto et al. (1995)
VTG in vitro, in vivo trout study
Soto et al. (1995), Matta et al. (1998)
RER (MCF-7-luc), E-screen,
CB-ER, in vivo—vaginal cell
cornification in mice
YES, E-screen RER (MCF-7-luc)
Joyeux et al. (1997), Soto et al. (1995),
Beard et al. (1999), Kramer et al. (1997)
Tran et al. (1996), Chaloupka et al. (1992),
Santodonato (1997), Clemons et al. (1998)
YES
Tran et al. (1996)
In vitro substrate stimulated
testosterone production by Leydig
cells
In vivo juvenile rainbow trout
exposure
In vivo rat feeding exposure
Laskeyand Phelps (1991)
Lafuente et al. (1997)
In vivo rat feeding study
Ronis et al. (1998)
Ricard et al. (1998)
J.P. Giesy et al. / Marine Pollution Bulletin 45 (2002) 3–16
9
Table 3
Examples of wild type and recombinant cell lines used for assessment of AhR- and ER-mediated activity
Cell line
Reporter gene
Species of
origin
Response
measured
Reference
EROD
ECOD
EROD
EROD
luc
Van Den Heuvel et al. (1994), Sanderson et al. (1996)
Hoivik et al. (1997)
Clemons et al. (1997)
Hahn et al. (1996), Nakama et al. (1997)
Garrison et al. (1996), Sanderson et al. (1996), Murk et al.
(1996)
Villeneuve et al. (1997)
Garrison et al. (1996), Anderson et al. (1999)
Garrison et al. (1996)
AhR
H4IIE
HepG2
RLT-W1
PLHC-1
H4IIE ¼ CALUX
luc
Rat
Human
Trout
Fish
Rat
HeLa
HepG2
GPC16
MLE-BV
MCF7, LS180
AHL
Hepa1
luc
luc
luc
luc
luc-transient
luc-transient
luc, luc-transient
Human
Human
Guinea pig
Mouse
Human
Hamster
Mouse
luc
luc
luc
Hepa1
HepG2, MCF7
H4IIE
GPC16
Hepal
RLT2.0
PAP-transient
PAP
PAP
luc
Mouse
Human
Rat
Guinea pig
Mouse
Trout
Clemons et al. (1998), Garrison et al. (1996), Balaguer et al.
(1996)
El-Fouly et al. (1994)
PAP
luc
El-Fouly et al. (1994)
Richter et al. (1997)
luc-transient
luc
luc
Human
Human
Human
luc
luc
luc
Nie et al. (2001), Clemons et al. (1998)
Kramer et al. (1997), Balaguer et al. (1996)
Legler et al. (1998)
ER
MCF-7
MCF-7 (MVLN)
T47D (ER-CALUX)
luc
Reporter genes: nothing, wild type cell line.
The character of the dose-response curves for endogenous enzyme activities controlled by the AhR
mechanism are biphasic with a decrease in response at
greater doses. Some chemicals inducing the cytochrome
P4501A1 activity can also serve as substrates for this
enzyme, so they cause competitive substrate inhibition
and reduced activity at greater concentrations (Hahn
et al., 1996; Garrison et al., 1996; Willett et al., 1997).
This problem is avoided in genetically engineered cell
lines, where the chemical inducers are not competitive substrates for the transfected reporter enzyme. Genetically engineered cells generally exhibit greater
sensitivity, dynamic range and selectivity than their
corresponding wild type cells (Sanderson et al., 1996;
Murk et al., 1996). Wild type and recombinant cell lines
have been developed mostly for mammal and teleost
species. Studies are being conducted to develop cell lines
for other species, including amphibians and reptiles.
Both immortalized (continuous) and primary cell cultures of primary hepatocytes from birds (Kennedy et al.,
1996a,b), mammals (Till et al., 1997) or fish have been
used to measure dioxin-like activity (Kennedy et al.,
1996a; Till et al., 1997) or xenoestrogenicity (Pelissero
et al., 1993). The responsiveness of assays as characterized by maximal fold induction relative to control,
sensitivity, detection limit and variability is species- and
cell line-specific. Differences among species and tissues
in ligand-binding affinity, ligand specificity and physicochemical properties of the receptor have been shown
along with significant differences in responsiveness to
standard ligands (El-Fouly et al., 1994). Observed
differences in responsiveness are explained by species
differences in the level and structure of the receptors and
their associated proteins, and/or transacting factors
present in each cell line (Joyeux et al., 1997; Garrison
et al., 1996). Studies comparing responsiveness among
cell lines from different species (mostly mammals and
fish) to single compounds or mixtures revealed substantial differences between the relative potencies
derived from different species (van den Berg et al., 1998;
Clemons et al., 1997). Also the time course of response
differs among cell lines. Fish cells have been observed to
be slower in responsiveness than the mammalian cells
(Richter et al., 1997). Therefore, the optimum duration
of exposure is important to obtain reproducible results
and is cell-line specific ranging from 6 to 72 h (Anderson
et al., 1999).
Both estrogenic and antiestrogenic effects can be assessed with ER-responsive cell lines. Antiestrogenicity
can be detected directly by growing cells in medium
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deficient in 17b-estradiol (E2 ) or by the antagonism of
co-administered E2 (Kramer et al., 1997). As one of the
important mechanisms of antiestrogenicity, modulation
of endocrine pathways by AhR agonists has been observed (Navas and Segner, 1998). TCDD and related
compounds have been observed to be antiestrogenic in
in vitro tests but also in some in vivo studies (Gillesby and
Zacharewski, 1998). The interactions of the TCDD- and
E2 -induced signaling pathways are complex; AhR agonists are antiestrogenic via direct interactions between
the nuclear AhR and genomic sequences in flanking regions of E2 -regulated genes (Safe and Krishnan, 1995).
Two-way cross-talk between the intracellular signaling
pathways involving AhR agonists and estrogens by
mutual inhibition of binding of ER or AhR to DNA has
been reported (Kharat and Saatcioglu, 1996), but recently this observation has been disputed by Hoivik et al.
(1997).
In extracts containing complex mixtures of compounds, potential cytotoxicity should be evaluated in
bioassays with the same cells that were used in receptor
mediated effects. This is because the cytotoxic effects
could mask potential antiestrogenic or other types of
effects. Other mechanisms of antiestrogenicity include:
(1) competitive binding of the ligand to the ER displacing E2 ; (2) increased E2 metabolism due to induction
of xenobiotic metabolizing enzymes; (3) inhibition of E2 induced gene expression; (4) ER down regulation (Gillesby and Zacharewski, 1998). Anti-/estrogenic potential
of some compounds changes depending on the E2 concentration (Kramer et al., 1997), thus testing only in
media without any E2 does not adequately assess the
physiological situation where there is always some E2
present.
1997) can be assessed after extraction. Paper mill effluent fractions elicited strong TCDD-like and antiestrogenic activity (Nie et al., 2001), whereas significant
estrogenic and TCDD-like activity has been detected
in crude extract of inhalable air particulate matter
(Clemons et al., 1998) or from diesel exhaust particles
(Meek, 1998). Significant dioxin-like activity has been
observed in eggs of birds such as herring gull, cormorant, and great blue heron (Tillitt et al., 1991; Kennedy
et al., 1996b) as well as in birds at different stages of
development (Jones et al., 1994). Among other animals,
extracts of fish (white sucker, juvenile whitefish) (Van
Den Heuvel et al., 1994; Koistinen et al., 1998) and otter
(Murk et al., 1998) have also been tested. For the biota
samples either whole body extracts or more specific
tissue extracts, especially livers have been used. An important step is the sample preparation and extraction.
Direct water sample or extracts prepared with organic
solvents can be used. Solid samples are usually extracted
by organic solvents. The solvent of choice needs to be
compatible with the cell system, not causing any effect
by itself, but enabling distribution of the extracted material to the cells.
Extracts can be cytotoxic, which is caused by some
compounds present in complex mixtures. For example,
sulfur is a major cytotoxic constituent in sediment extracts, which should be eliminated prior to performing
dioxin-like or estrogenic activities. The measurement of
cell viability/cytotoxicity is essential in all bioassays
dealing with complex mixtures as well as single compounds. Cell bioassays with 96 well plates enable the
measurement of several samples at the same time. In
addition, current procedures allow subsequent measurement of viability index, enzyme activity and protein
content in the same 96 well plates (Blankenship et al.,
2000).
5. Testing of complex mixtures with bioassays
In vitro bioassays have been used to assess TCDDlike and estrogenic activity in a variety of environmental
matrices, both abiotic and biotic (Khim et al., 1999;
Kannan et al., 2000; Hilscherova et al., 2000). Various
aquatic samples, such as porewater (Nakama et al.,
1997), stream water (Villeneuve et al., 1997), extracts
from waste water treatment plant influent and effluents,
sediments (Murk et al., 1996) or settling particulate
matter (Balaguer et al., 1996; Pons et al., 1990; Koistinen et al., 1998; Engwall et al., 1997; Brunstrom et al.,
1992) have been analyzed by in vitro cell bioassays.
Extracts from semi-permeable membrane devices enabled examination of concentrations of in situ bioavailable lipophilic contaminants to which aquatic
organisms are exposed (Villeneuve et al., 1997). Also
sludge (Legler et al., 1998; Koistinen et al., 1998) or
atmospheric samples including air particulates (Clemons
et al., 1998) and fly ash from incinerators (Till et al.,
6. Estimation of relative potencies of complex mixtures
The relative potencies of samples are usually calculated as the amount of standard (reference toxicant)
giving the same response as the sample, commonly
based on the amount of sample needed to produce 50%
of the maximal standard response (EC50 ). The exogenous compound with the greatest known affinity as well
as toxicity, TCDD, is used as a standard for AhRmediated responses. The endogenous substrate E2 serves
as a standard for ER-mediated activity. Activities of
samples are then expressed as bioassay-derived equivalents: dioxin equivalents (TCDD-EQ) or estradiol
equivalents (E2 -EQ) per specified amount of sample.
For calculating and comparing the equivalents complete
dose-response curves from step-wise diluted extracts and
standards should be obtained. This is rather difficult
with complex extracts. Common problems encountered
J.P. Giesy et al. / Marine Pollution Bulletin 45 (2002) 3–16
when determining the relative potencies of complex
mixtures include different efficacy (maximal induced response), non-parallel slopes, cytotoxicity at greater
concentrations or insufficient mass of agonists to reach
full efficacy or the occurrence of partial agonists that do
not attain the maximum possible response. These limitations must be taken into account when calculating the
relative potency (RP) of the sample. There is always
variation in the EC50 in replicates measured on different
days due to differences in plating density of cells. For
some cell lines the normalization for protein content can
solve this problem. For endogenous enzyme activities
the normalization to protein content is necessary. In
some transgenic cell lines the normalization to the
amount of protein present has been inadvisable because
of increased variability of the normalized results. Protein normalization is not recommended in cell lines used
for estrogen-receptor mediated activity, where response
induction correlates with estrogen-induced protein synthesis (Villeneuve et al., 1998).
Some non-active parent compounds can be metabolically activated to potent inducers of receptor-mediated response; alternatively the active compound can be
biotransformed to non-active metabolites. For most
compounds, the activity of their metabolites is unknown. Some of the cell lines possess a number of
metabolic capabilities and upon prolonged duration of
exposure they can partly simulate in vivo biotransformation of some compounds. This fact can be used analytically by use of selective inhibitors.
7. Potency-balance calculations
In the potency-balance approach, total activities determined by a bioassay are compared with the sum of
the potency of the individual compounds determined by
chemical analysis. This strategy has been widely used
for dioxin-like compounds (van den Berg et al., 1998;
Ahlborg et al., 1992) and for estrogenic compounds
(Safe, 1995a). Toxic equivalents (TEQs) are calculated
by multiplying the RP for the specific assay (if available)
or the international toxic equivalency factor by concentration of the specific congener giving total sum
TEQs per mass unit. For calculating the TEQs from
chemical data effects are assumed to be additive (Eq. (1)).
TEFs are species-, endpoint- and assay specific determination of potency expressed relative to the standard, they can vary widely depending on the species
and endpoint. The RPs should be used for bioassaydirected potency-balance calculation for complex mixtures,
TEQ ¼
N
X
i¼1
Conc: of compoundi TEFi
ð1Þ
11
they are specific for studied endpoint and assay (Villeneuve et al., 1999).
The international dioxin TEFs are consensus values,
based on many different types of assays (van den Berg
et al., 1998) including multiple in vitro and in vivo
endpoints for multiple species. TEF values are orderof-magnitude estimates suitable for risk assessment
purposes. Because of the differences in RPs among
species, specific sets of international TEFs have been
established for mammals, fish and birds (van den Berg
et al., 1998). Currently TEFs and RPs are available for
dioxins, furans, some PCBs and PAHs from a number
of assays. There are many compounds with potential
AhR-mediated activity for which RPs are unavailable
and TEFs have not been established (Villeneuve et al.,
2000). Therefore those compounds cannot be included
in the potency-balance calculations.
Limited data are available for the RPs of estrogenic
compounds; RPs have been established only by use of
in vitro bioassays for a few alkylphenolic compounds
and PAHs (Villeneuve et al., 1998; Clemons et al., 1998).
In this case by calculating the E2 -EQs based on analytical results one can estimate the proportion of the
total activity determined by bioassay that is represented
by the compounds which have been quantified and have
known relative potencies. There are several limitations
of calculating TEQs from analytical results: (1) RPs or
TEFs are available for only limited number of chemicals. For some compounds there are no endpoint-specific or consensus values for TEFs available; (2) the use
of TEFs derived for other species, usually from mammals, where the most research has been conducted, to
non-mammalian species may not be suitable due to the
interspecies differences in sensitivity; (3) there may be
some compounds not routinely detected whose contribution to the activity would be overlooked; (4) application of the additive approach is routinely used in
the total activity calculation; potential interactions
among compounds in a mixture, such as synergism or
antagonism are neglected; (5) detailed analysis of trace
contaminants require specialized equipment such as
HRGC/HRMS (high resolution gas chromatograph/
mass spectrophotometer), which is not available in all
laboratories and may be prohibitively expensive. TEQs
estimated based on analytical data are correlated with
the bioassay results in some situations, depending on
the composition of the complex mixture of compounds
in the samples. For biota samples for which we report
here, highly significant correlations have been found
between bioassay derived EROD activity and instrumentally measured TEQs in extracts of fish or bird
samples (Van Den Heuvel et al., 1994; Tillitt et al., 1991;
Kennedy et al., 1996b). However, toxic activities determined in the bioassays and concentrations of known
dioxin-like or xenoestrogenic compounds are sometimes not correlated. For instance, data obtained from
12
J.P. Giesy et al. / Marine Pollution Bulletin 45 (2002) 3–16
bioassays may be an independent parameter that is
predictive of ecotoxicological effects. Besides non-additive (synergistic or antagonistic) interactions among
individual ligands, differences between TEQs derived in
bioassays and those calculated from concentrations of
individual compounds may be caused by the following
events: (1) there are some other active compounds present, which were not identified by the chemical analysis
(Willett et al., 1997); (2) non-complete dose responses or
cytotoxicity disabling accurate estimations of TEQs; (3)
the RPs or TEFs used may not be appropriate. Generally, bioassay data have great ecotoxicological relevance because they represent an integrated biological
response.
It is necessary to point out disadvantages and limitations of in vitro bioassays. Bioassays do not account
for the pharmacokinetics, tissue distribution and biotransformation that may occur in vivo. If cell lines
possess only limited metabolic activities, substances active after bioactivation may not be detected by in vitro
system (Villeneuve et al., 1997). Bioassays do not identify the individual compounds causing the response.
Bioassays assess only the activity of compounds that act
through a specific receptor-mediated mechanism of
action. The non-receptor-mediated responses, such as
estrogen-like chemicals acting through different mechanisms, are not taken into consideration.
8. Fractionation approach
In vitro bioassays can be used in combination with
specific analytical techniques as a bioassay-directed
fractionation methodology. This approach provides information needed for monitoring and risk assessment of
the compounds with specific modes of action and may
lead to identification of novel classes of environmental
toxicants (Brunstrom et al., 1992). If complex mixtures
cause a significant response in a bioassay in order to
determine the causes and identify possible sources, the
compounds causing the observed response need to be
identified. Instrumental analysis could be applied to the
entire mixture or sub-fractions. Recommended strategy
for toxicants identification and evaluation in complex
mixtures is shown (Fig. 2). The general steps are: (1)
screening of the whole extract—to determine the samples containing significant toxic potencies, which require
further chemical analysis. If no significant response is
observed, there is no need to conduct expensive, timeand material-consuming chemical analysis. Since the
method detection limit is known for the bioassay, an
upper limit of concentration of TEQs in the sample can
be defined; (2) fractionation of the samples that were
active in bioassays and chromatographic analysis can be
used to determine the most probable contributors to the
total activity; (3) generating the full dose-response re-
Fig. 2. Screening system: Toxicant identification and evaluation
strategy.
lationship of the unfractionated sample or fractions
thereof, so that the total activity of the sample can be
determined as response equivalents. Calculation of the
potency balance is accomplished by comparing the activity observed in the bioassay with the potential activity
based identification and quantification by instrumental
analyses. If the values derived and fractionation do not
indicate that there were antagonistic effects in the whole
extract, it can be concluded that all of the significant
contributors to the total complex mixture have been
identified. However, if the total activity determined
from the bioassay is significantly greater than those
predicted from the instrumental it can be inferred that
there are unidentified compounds or that there is synergism. Again by comparing the activity of the whole
extract to that of the various fractions, it is possible to
determine if the difference is due to the presence of
unidentified compounds or synergism. In our studies,
we have found that antagonisms can occur, particularly
between non-AhR-active and AhR-active PCB congeners.
J.P. Giesy et al. / Marine Pollution Bulletin 45 (2002) 3–16
To apply the potency-balance approach with complex
mixtures, species- and endpoint-specific RPs/TEFs and
especially E2 -EQs need to be determined. Fractionation
of whole extracts into groups of compounds with similar
characteristics and subsequent bioassay testing can be
useful in determining the most appropriate instrumental
analysis that should be applied and can prevent application of non-essential and costly analysis of the fractions with low activity and thus significance (Engwall
et al., 1997). For most compounds, fractionation based
on polarity and/or molecular size of the compounds is
generally suitable. These characteristics are easily selected for with simple chromatographic techniques. Instrumental analyses can be applied to determine the
compounds responsible for the activity observed in each
fraction. For instance, if the activity was observed in a
more polar fraction normal phase liquid chromatography might be deemed more appropriate than gas
chromatography, or derivatisation might be deemed
appropriate before subsequent analyses.
9. Conclusions
Many studies have demonstrated the utility of bioassays in assessment of receptor-mediated activities of
both individual chemicals and complex mixtures. Bioassays can be used for the detection and quantitation of
receptor agonists/antagonists in complex mixtures, thus
providing a relative measure of bioactive compounds in
food, biological, or abiotic samples. Bioassays can also
be useful for identification and characterization of novel
receptor agonists, for examination of species differences
in receptor-mediated responses or effectiveness of remediation procedures designed to decrease specific type of
contamination. Bioassays are also useful screening tools
for identifying responsible compounds following fractionation of a complex mixture, they enable to prioritize
samples which require further investigation. In vitro cell
bioassays are excellent systems for evaluating the activities of chemicals with specific mode of action. Bioassays, based on in vitro responses of cells, including
both wild type or recombinant (genetically modified) cell
lines can also be used for assessment of other toxicologically and pharmacologically important chemicals
where ligand-dependent induction of gene expression
has been demonstrated. Such compounds include xenoandrogens, heavy metals and compounds that can
cause induction of peroxisome proliferation.
Acknowledgements
Preparation of the manuscript as well as the research
on which it is based was supported by the Czech Ministry of Education (CEZ: J07/98:1410003) and Ministry
13
of Agriculture (MZE-M03-99-01). A Fullbright fellowship to K. Hilscherova is gratefully acknowledged.
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